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Year: 2005

Effects of biodiversity on ecosystem functioning: a consensus of

current knowledge

Hooper, D U; Chapin III, F S; Ewel, J J; Hector, A; Inchausti, P; Lavorel, S; Lawton, J

H; Lodge, D M; Loreau, M; Naeem, S; Schmid, B; Setälä, H; Symstad, A J;

Vandermeer, J; Wardle, D A

Hooper, D U; Chapin III, F S; Ewel, J J; Hector, A; Inchausti, P; Lavorel, S; Lawton, J H; Lodge, D M; Loreau, M;Naeem, S; Schmid, B; Setälä, H; Symstad, A J; Vandermeer, J; Wardle, D A. Effects of biodiversity on ecosystemfunctioning: a consensus of current knowledge. Ecological Monographs 2005, 75(1):3-35.Postprint available at:http://www.zora.unizh.ch

Posted at the Zurich Open Repository and Archive, University of Zurich.http://www.zora.unizh.ch

Originally published at:Ecological Monographs 2005, 75(1):3-35

Hooper, D U; Chapin III, F S; Ewel, J J; Hector, A; Inchausti, P; Lavorel, S; Lawton, J H; Lodge, D M; Loreau, M;Naeem, S; Schmid, B; Setälä, H; Symstad, A J; Vandermeer, J; Wardle, D A. Effects of biodiversity on ecosystemfunctioning: a consensus of current knowledge. Ecological Monographs 2005, 75(1):3-35.Postprint available at:http://www.zora.unizh.ch

Posted at the Zurich Open Repository and Archive, University of Zurich.http://www.zora.unizh.ch

Originally published at:Ecological Monographs 2005, 75(1):3-35

Effects of biodiversity on ecosystem functioning: a consensus of

current knowledge

Abstract

Humans are altering the composition of biological communities through a variety of activities thatincrease rates of species invasions and species extinctions, at all scales, from local to global. Thesechanges in components of the Earth's biodiversity cause concern for ethical and aesthetic reasons, butthey also have a strong potential to alter ecosystem properties and the goods and services they provide tohumanity. Ecological experiments, observations, and theoretical developments show that ecosystemproperties depend greatly on biodiversity in terms of the functional characteristics of organisms presentin the ecosystem and the distribution and abundance of those organisms over space and time. Specieseffects act in concert with the effects of climate, resource availability, and disturbance regimes ininfluencing ecosystem properties. Human activities can modify all of the above factors; here we focuson modification of these biotic controls.

The scientific community has come to a broad consensus on many aspects of the relationship betweenbiodiversity and ecosystem functioning, including many points relevant to management of ecosystems.Further progress will require integration of knowledge about biotic and abiotic controls on ecosystemproperties, how ecological communities are structured, and the forces driving species extinctions andinvasions. To strengthen links to policy and management, we also need to integrate our ecologicalknowledge with understanding of the social and economic constraints of potential managementpractices. Understanding this complexity, while taking strong steps to minimize current losses ofspecies, is necessary for responsible management of Earth's ecosystems and the diverse biota theycontain.

Based on our review of the scientific literature, we are certain of the following conclusions:

1)Species' functional characteristics strongly influence ecosystem properties. Functional characteristicsoperate in a variety of contexts, including effects of dominant species, keystone species, ecologicalengineers, and interactions among species (e.g., competition, facilitation, mutualism, disease, andpredation). Relative abundance alone is not always a good predictor of the ecosystem-level importanceof a species, as even relatively rare species (e.g., a keystone predator) can strongly influence pathwaysof energy and material flows.

2)Alteration of biota in ecosystems via species invasions and extinctions caused by human activities hasaltered ecosystem goods and services in many well-documented cases. Many of these changes aredifficult, expensive, or impossible to reverse or fix with technological solutions.

3)The effects of species loss or changes in composition, and the mechanisms by which the effectsmanifest themselves, can differ among ecosystem properties, ecosystem types, and pathways ofpotential community change.

4)Some ecosystem properties are initially insensitive to species loss because (a) ecosystems may havemultiple species that carry out similar functional roles, (b) some species may contribute relatively littleto ecosystem properties, or (c) properties may be primarily controlled by abiotic environmentalconditions.

5)More species are needed to insure a stable supply of ecosystem goods and services as spatial andtemporal variability increases, which typically occurs as longer time periods and larger areas areconsidered.

We have high confidence in the following conclusions:

1)Certain combinations of species are complementary in their patterns of resource use and can increaseaverage rates of productivity and nutrient retention. At the same time, environmental conditions caninfluence the importance of complementarity in structuring communities. Identification of which andhow many species act in a complementary way in complex communities is just beginning.

2)Susceptibility to invasion by exotic species is strongly influenced by species composition and, undersimilar environmental conditions, generally decreases with increasing species richness. However,several other factors, such as propagule pressure, disturbance regime, and resource availability alsostrongly influence invasion success and often override effects of species richness in comparisons acrossdifferent sites or ecosystems.

3)Having a range of species that respond differently to different environmental perturbations canstabilize ecosystem process rates in response to disturbances and variation in abiotic conditions. Usingpractices that maintain a diversity of organisms of different functional effect and functional responsetypes will help preserve a range of management options.

Uncertainties remain and further research is necessary in the following areas:

1)Further resolution of the relationships among taxonomic diversity, functional diversity, andcommunity structure is important for identifying mechanisms of biodiversity effects.

2)Multiple trophic levels are common to ecosystems but have been understudied inbiodiversity/ecosystem functioning research. The response of ecosystem properties to varyingcomposition and diversity of consumer organisms is much more complex than responses seen inexperiments that vary only the diversity of primary producers.

3)Theoretical work on stability has outpaced experimental work, especially field research. We needlong-term experiments to be able to assess temporal stability, as well as experimental perturbations toassess response to and recovery from a variety of disturbances. Design and analysis of such experimentsmust account for several factors that covary with species diversity.

4)Because biodiversity both responds to and influences ecosystem properties, understanding thefeedbacks involved is necessary to integrate results from experimental communities with patterns seenat broader scales. Likely patterns of extinction and invasion need to be linked to different drivers ofglobal change, the forces that structure communities, and controls on ecosystem properties for thedevelopment of effective management and conservation strategies.

5)This paper focuses primarily on terrestrial systems, with some coverage of freshwater systems,because that is where most empirical and theoretical study has focused. While the fundamentalprinciples described here should apply to marine systems, further study of that realm is necessary.

Despite some uncertainties about the mechanisms and circumstances under which diversity influencesecosystem properties, incorporating diversity effects into policy and management is essential, especiallyin making decisions involving large temporal and spatial scales. Sacrificing those aspects of ecosystemsthat are difficult or impossible to reconstruct, such as diversity, simply because we are not yet certainabout the extent and mechanisms by which they affect ecosystem properties, will restrict futuremanagement options even further. It is incumbent upon ecologists to communicate this need, and thevalues that can derive from such a perspective, to those charged with economic and policydecision-making.

3

Ecological Monographs, 75(1), 2005, pp. 3–35q 2005 by the Ecological Society of America

ESA Report

EFFECTS OF BIODIVERSITY ON ECOSYSTEM FUNCTIONING:A CONSENSUS OF CURRENT KNOWLEDGE

D. U. HOOPER,1,16 F. S. CHAPIN, III,2 J. J. EWEL,3 A. HECTOR,4 P. INCHAUSTI,5 S. LAVOREL,6 J. H. LAWTON,7

D. M. LODGE,8 M. LOREAU,9 S. NAEEM,10 B. SCHMID,4 H. SETALA,11 A. J. SYMSTAD,12

J. VANDERMEER,13 AND D. A. WARDLE14,15

1Department of Biology, Western Washington University, Bellingham, Washington 98225 USA2Institute of Arctic Biology, University of Alaska, Fairbanks, Alaska 99775 USA

3Institute of Pacific Islands Forestry, Pacific Southwest Research Station, USDA Forest Service, 1151 Punchbowl Street,Room 323, Honolulu, Hawaii 96813 USA

4Institute of Environmental Sciences, University of Zurich, Winterthurerstrasse 190, CH-8057 Zurich, Switzerland5CEBC-CNRS, 79360 Beauvoir-sur-Niort, France

6Laboratoire d’Ecologie Alpine, CNRS UMR 5553, Universite J. Fourier, BP 53, 38041 Grenoble Cedex 9, France7Natural Environment Research Council, Polaris House, North Star Avenue, Swindon SN2 1EU, UK

8Department of Biological Sciences, P.O. Box 369, University of Notre Dame, Notre Dame, Indiana 46556-0369 USA9Laboratoire d’Ecologie, UMR 7625, Ecole Normale Superieure, 46 rue d’Ulm, 75230 Paris Cedex 05, France

10Department of Ecology, Evolution and Environmental Biology, Columbia University, 1200 Amsterdam Avenue,New York, New York 10027 USA

11University of Helsinki, Department of Ecological and Environmental Sciences, Niemenkatu 73, FIN-15140 Lahti, Finland12U.S. Geological Survey, Mount Rushmore National Memorial, 13000 Highway 244, Keystone, South Dakota 57751 USA

13Department of Biology, University of Michigan, Ann Arbor, Michigan 48109 USA14Landcare Research, P.O. Box 69, Lincoln, New Zealand

15Department of Forest Vegetation Ecology, Swedish University of Agricultural Sciences, SE901-83, Umea, Sweden

Abstract. Humans are altering the composition of biological communities through avariety of activities that increase rates of species invasions and species extinctions, at allscales, from local to global. These changes in components of the Earth’s biodiversity causeconcern for ethical and aesthetic reasons, but they also have a strong potential to alterecosystem properties and the goods and services they provide to humanity. Ecologicalexperiments, observations, and theoretical developments show that ecosystem propertiesdepend greatly on biodiversity in terms of the functional characteristics of organisms presentin the ecosystem and the distribution and abundance of those organisms over space andtime. Species effects act in concert with the effects of climate, resource availability, anddisturbance regimes in influencing ecosystem properties. Human activities can modify allof the above factors; here we focus on modification of these biotic controls.

The scientific community has come to a broad consensus on many aspects of the re-lationship between biodiversity and ecosystem functioning, including many points relevantto management of ecosystems. Further progress will require integration of knowledge aboutbiotic and abiotic controls on ecosystem properties, how ecological communities are struc-tured, and the forces driving species extinctions and invasions. To strengthen links to policyand management, we also need to integrate our ecological knowledge with understandingof the social and economic constraints of potential management practices. Understandingthis complexity, while taking strong steps to minimize current losses of species, is necessaryfor responsible management of Earth’s ecosystems and the diverse biota they contain.

Manuscript received 2 June 2004; accepted 7 June 2004; final version received 7 July 2004. Corresponding Editor (ad hoc): J. S.Denslow. This article is a committee report commissioned by the Governing Board of the Ecological Society of America. Reprintsof this 33-page ESA report are available for $5.00 each, either as pdf files or as hard copy. Prepayment is required. Order reprintsfrom the Ecological Society of America, Attention: Reprint Department, 1707 H Street, N.W., Suite 400, Washington, DC 20006USA (e-mail: [email protected]).

16 E-mail: [email protected]

4 ESA REPORT Ecological MonographsVol. 75, No. 1

Based on our review of the scientific literature, we are certain of the following con-clusions:

1) Species’ functional characteristics strongly influence ecosystem properties. Func-tional characteristics operate in a variety of contexts, including effects of dominant species,keystone species, ecological engineers, and interactions among species (e.g., competition,facilitation, mutualism, disease, and predation). Relative abundance alone is not always agood predictor of the ecosystem-level importance of a species, as even relatively rare species(e.g., a keystone predator) can strongly influence pathways of energy and material flows.

2) Alteration of biota in ecosystems via species invasions and extinctions caused byhuman activities has altered ecosystem goods and services in many well-documented cases.Many of these changes are difficult, expensive, or impossible to reverse or fix with tech-nological solutions.

3) The effects of species loss or changes in composition, and the mechanisms by whichthe effects manifest themselves, can differ among ecosystem properties, ecosystem types,and pathways of potential community change.

4) Some ecosystem properties are initially insensitive to species loss because (a) eco-systems may have multiple species that carry out similar functional roles, (b) some speciesmay contribute relatively little to ecosystem properties, or (c) properties may be primarilycontrolled by abiotic environmental conditions.

5) More species are needed to insure a stable supply of ecosystem goods and servicesas spatial and temporal variability increases, which typically occurs as longer time periodsand larger areas are considered.

We have high confidence in the following conclusions:

1) Certain combinations of species are complementary in their patterns of resource useand can increase average rates of productivity and nutrient retention. At the same time,environmental conditions can influence the importance of complementarity in structuringcommunities. Identification of which and how many species act in a complementary wayin complex communities is just beginning.

2) Susceptibility to invasion by exotic species is strongly influenced by species com-position and, under similar environmental conditions, generally decreases with increasingspecies richness. However, several other factors, such as propagule pressure, disturbanceregime, and resource availability also strongly influence invasion success and often overrideeffects of species richness in comparisons across different sites or ecosystems.

3) Having a range of species that respond differently to different environmental perturbationscan stabilize ecosystem process rates in response to disturbances and variation in abiotic con-ditions. Using practices that maintain a diversity of organisms of different functional effect andfunctional response types will help preserve a range of management options.

Uncertainties remain and further research is necessary in the following areas:

1) Further resolution of the relationships among taxonomic diversity, functional diversity,and community structure is important for identifying mechanisms of biodiversity effects.

2) Multiple trophic levels are common to ecosystems but have been understudied inbiodiversity/ecosystem functioning research. The response of ecosystem properties to vary-ing composition and diversity of consumer organisms is much more complex than responsesseen in experiments that vary only the diversity of primary producers.

3) Theoretical work on stability has outpaced experimental work, especially field re-search. We need long-term experiments to be able to assess temporal stability, as well asexperimental perturbations to assess response to and recovery from a variety of disturbances.Design and analysis of such experiments must account for several factors that covary withspecies diversity.

4) Because biodiversity both responds to and influences ecosystem properties, under-standing the feedbacks involved is necessary to integrate results from experimental com-munities with patterns seen at broader scales. Likely patterns of extinction and invasionneed to be linked to different drivers of global change, the forces that structure communities,and controls on ecosystem properties for the development of effective management andconservation strategies.

5) This paper focuses primarily on terrestrial systems, with some coverage of freshwatersystems, because that is where most empirical and theoretical study has focused. While the

February 2005 5ECOSYSTEM EFFECTS OF BIODIVERSITY

fundamental principles described here should apply to marine systems, further study of thatrealm is necessary.

Despite some uncertainties about the mechanisms and circumstances under which di-versity influences ecosystem properties, incorporating diversity effects into policy andmanagement is essential, especially in making decisions involving large temporal and spatialscales. Sacrificing those aspects of ecosystems that are difficult or impossible to reconstruct,such as diversity, simply because we are not yet certain about the extent and mechanismsby which they affect ecosystem properties, will restrict future management options evenfurther. It is incumbent upon ecologists to communicate this need, and the values that canderive from such a perspective, to those charged with economic and policy decision-making.

Key words: biodiversity; complementary resource use; ecosystem goods and services; ecosystemprocesses; ecosystem properties; functional characteristics; functional diversity; net primary produc-tion; sampling effect; species extinction; species invasions; species richness; stability.

I. INTRODUCTION

A. The context: human effects on biodiversity

Human activities have been and are continuing tochange the environment on local and global scales.Many of these alterations are leading to dramaticchanges in the biotic structure and composition of eco-logical communities, either from the loss of species orfrom the introduction of exotic species. Such changescan readily change the ways in which ecosystems work.Altered biodiversity has led to widespread concern fora number of both market (e.g., ecotourism, ‘‘mining’’for medicines) and non-market (e.g., ethical, aesthetic)reasons (Barbier et al. 1995, Kunin and Lawton 1996,Schwartz et al. 2000, Hector et al. 2001b, Minns et al.2001, Sax and Gaines 2003). These reasons are com-pelling in their own right, but ecologists have raisedadditional concerns: What is the effect of changingbiodiversity on ecosystem properties, such as produc-tivity, carbon storage, hydrology, and nutrient cycling?The obvious follow-up question is: What are the con-sequences of such largely anthropogenic changes inbiodiversity on the goods and services that ecosystemsprovide to humans? If altered biodiversity affects eco-system properties, is there a point at which changes inproperties might adversely influence human welfare?

While global extinction of a species is clearly animportant conservation concern, local species extinc-tions or even large changes in abundances have as muchpotential to affect ecosystem properties (e.g., Zimov etal. 1995). Local extinctions and large effects of intro-duced species are more common than global extinctionsand can be very difficult to reverse, as seen with manyattempts to reintroduce species or eradicate invasiveexotics (Enserink 1999, Finkel 1999, Kaiser 1999, Ma-lakoff 1999, Stokstad 1999, Stone 1999, Sax andGaines 2003). These problems affect both managed andunmanaged ecosystems (Pimentel et al. 1992).

The effects of biodiversity loss or changes in com-munity composition on the functioning of ecosystemshave been the focus of much ecological research, with

an explosion of research over the past decade (Schulzeand Mooney 1993, Kinzig et al. 2002, Loreau et al.2002b). In spite of this effort, however, there remainimportant aspects that are still not well understood.There has been substantial debate within the ecologicalcommunity on the interpretation of some recent re-search and whether the findings from these studies areas important as other factors that are well known tocorrelate with ecosystem functioning in nature. Manyof the authors of this paper have been on different sidesof this debate. Our goals here are to summarize a con-sensus view for the ecological community of currentunderstandings of the relationships between biodiver-sity and ecosystem functioning with an eye to uncer-tainties and future directions that can help to addresssome of these uncertainties. We review the scientificevidence for links between biodiversity and ecosystemfunctioning, including theoretical, observational, andexperimental results, and we link the scientific studiesto potential management and policy implications. Weparticularly focus on research over the past decade thattreats quantitative aspects of biodiversity, since earlierwork has been summarized elsewhere (Schulze andMooney 1993). We highlight areas of consensus amongecologists, point out areas of disagreement, and suggestquestions for future study.

B. Definitions

Clear discussion of the effects of biodiversity onecosystem functioning requires clear definitions ofthese two terms. The term biodiversity encompasses abroad spectrum of biotic scales, from genetic variationwithin species to biome distribution on the planet (Wil-son 1992, Gaston 1996, Purvis and Hector 2000, Moo-ney 2002). Biodiversity can be described in terms ofnumbers of entities (how many genotypes, species, orecosystems), the evenness of their distribution, the dif-ferences in their functional traits (Box 1), and theirinteractions. While biodiversity has often been used asa synonym for species richness (the number of speciespresent), different components of biodiversity (e.g.,


BOX 1. Functional traits, functional types, and functional diversity

An understanding of how changes in species richness and composition, and biodiversity in general, influenceecosystem properties requires an understanding of the functional traits of the species involved. By definition, functionaltraits are those that influence ecosystem properties or species’ responses to environmental conditions. Species areoften grouped together according to their functional traits to understand general mechanisms or to make studies ofcomplex systems more tractable. Functional types (aka functional groups) are, at first glance, a relatively simpleconcept. A functional type is a set of species that have similar effects on a specific ecosystem process or similarresponses to environmental conditions. Functional types are similar to the guild concept from animal communityecology (Root 1967, Simberloff and Dayan 1991, Wilson 1999) and to niche concepts (Leibold 1995). Althoughfunctional types can be quite useful, the practice of defining them and quantifying functional diversity can be difficult.There are four basic reasons for this:

1) Organisms’ effects on ecosystem properties generally fall along a continuous gradient, not into distinct groups.Thus, designating functional groups may require arbitrary decisions as to where boundaries between groups lie. Inthe main text we use the term ‘‘functional types’’ to emphasize the functional axes differentiating species, rather thantheir specific groupings. Attention is now being directed towards alternative methods of quantifying both the diversityof functional traits of organisms and their effects on ecosystem properties (e.g., Grime et al. 1997b, Walker et al.1999, Lavorel and Garnier 2001, Petchey 2002).

2) Traits that determine how a species responds to a disturbance or change in environment (functional responsetraits) may differ from those that determine how that species affects ecosystem properties (functional effect traits;Lavorel et al. 1997, Landsberg 1999, Walker et al. 1999, Lavorel and Garnier 2002). Recent studies on biodiversity/ecosystem functioning have focused primarily on functional effect traits (Hooper and Vitousek 1997, Tilman et al.1997a, Hooper and Vitousek 1998, Emmerson et al. 2001). Studies of how species distributions may change in responseto climate change have focused primarily on functional response traits (e.g., Box 1996, Steffen 1996, Cramer 1997,Smith et al. 1997, Elmqvist et al. 2003). Response and effect traits may or may not be correlated with one another(Chapin et al. 1996a, Lavorel and Garnier 2002). Understanding links among functional response and effect traitsremains a significant challenge, but is critical to understanding the dynamics of ecosystem functioning in a changingworld (Lavorel and Garnier 2001, Hooper et al. 2002).

3) Functional types identified for a specific ecosystem property are not necessarily relevant to other properties.Defining types based on just a few traits known to affect many functions (such as specific leaf area, plant height, andseed mass; Westoby 1998, Grime 2001) may alleviate this problem, but whether such types yield insights into bio-diversity/ecosystem functioning relationships within ecosystems remains unknown.

4) Is functional diversity correlated with species diversity in natural ecosystems? The answer to this question dependsin part on mechanisms of community assembly (Fridley 2001, Hooper et al. 2002, Mouquet et al. 2002). The conceptsof niche differentiation and limiting similarity imply that functional characteristics of coexisting organisms must differat some level, which means that increasing species richness should lead to increasing functional diversity (Bazzaz1987, Weiher and Keddy 1999a, Dıaz and Cabido 2001, Schmid et al. 2002b). On the other hand, strong environmentalfilters could limit species composition to a relatively restricted range of functional characteristics (Pearson and Ro-senberg 1978, Dıaz et al. 1998, Weiher and Keddy 1999a, Dıaz and Cabido 2001, Loreau et al. 2001, Lavorel andGarnier 2002), thereby limiting the degree of functional diversity capable of influencing different ecosystem properties(Grime 2001). Increasing species richness would then just lead to finer division of the available niche space ratherthan to greater functional diversity (Dıaz and Cabido 2001, Enquist et al. 2002, Schmid et al. 2002b). Merging ourunderstanding of ecosystem level controls with our understanding of community dynamics and assembly is an importantfocus of future study (Thompson et al. 2001).

richness, relative abundance, composition, presence/absence of key species) can have different effects onecosystem properties. We are explicit in our use ofterminology in this paper, referring, for example, to‘‘species richness’’ when discussing numbers of spe-cies, ‘‘diversity’’ when discussing more general attri-butes including differences in relative abundance andcomposition, and ‘‘biodiversity’’ only when the broad-est scope of the term is warranted. In this paper, wefocus mostly on changes in richness and compositionat the species and functional type levels, not becausethey are always the most important, but because thatis where most research has concentrated. Effects ofgenetic and functional diversity within species, inter-

actions among species, and ecosystem diversity acrosslandscapes are areas that deserve greater attention.

The total suite of functional traits in a communityis one of the main determinants of ecosystem properties(Chapin et al. 1997, Chapin et al. 2000). We thereforediscuss the effects of biodiversity with respect to thefunctional traits of the species involved (see Box 1 andSection I.C., below). We do so in the context of gainor loss of species from a given site or ecosystem, ratherthan in terms of cross-system comparisons of diversitywhere other environmental variables are also chang-ing—though merging these perspectives begs for fur-ther study (see Sections I.C. and II.C., below). Thenumber of species alone may not be the best predictor


of ecosystem properties, and the relationship betweenspecies or taxonomic richness and functional diversityin natural ecosystems is still being explored (Dıaz andCabido 2001, Enquist et al. 2002, Hooper et al. 2002,Petchey 2002, Schmid et al. 2002b, Tilman et al. 2002;see also Box 1 and Section II.C.2).

Ecosystem functioning is also a broad term that en-compasses a variety of phenomena, including ecosys-tem properties, ecosystem goods, and ecosystem ser-vices (Christensen et al. 1996), although some re-searchers use the term ‘‘ecosystem functioning’’ assynonymous with ecosystem properties alone, exclu-sive of ecosystem goods and services. Ecosystem prop-erties include both sizes of compartments (e.g., poolsof materials such as carbon or organic matter) and ratesof processes (e.g., fluxes of materials and energy amongcompartments). Ecosystem goods are those ecosystemproperties that have direct market value. They includefood, construction materials, medicines, wild types fordomestic plant and animal breeding, genes for geneproducts in biotechnology, tourism, and recreation.Ecosystem services are those properties of ecosystemsthat either directly or indirectly benefit human endeav-ors, such as maintaining hydrologic cycles, regulatingclimate, cleansing air and water, maintaining atmo-spheric composition, pollination, soil genesis, and stor-ing and cycling of nutrients (Christensen et al. 1996,Daily 1997). Ecosystem properties vary among eco-systems, but levels, rates, or amounts of variability ofthese properties are not inherently ‘‘good’’ or ‘‘bad.’’This is in contrast to ecosystem goods and services, towhich humans attach value (although in some cases,the distinction between properties and services is notclear-cut). We refer to ‘‘ecosystem properties’’ to sum-marize the various pools and fluxes and to ‘‘ecosystemgoods and services’’ only when referring to the subsetof functioning of utilitarian value to humans.

When discussing effects of biodiversity on ecosys-tem functioning it is important to be specific aboutwhich components of biodiversity are affecting whichcomponents of functioning. Measures of process ratesand pool sizes include both levels (e.g., average ratesor sizes) and variation (amount of fluctuation). Varia-tion in ecosystem properties can result from fluctua-tions in the environment from year to year, directionalchanges in conditions, abiotic disturbance, or bioticdisturbance. There is no a priori reason to expect thatdifferent ecosystem properties have a single pattern ofresponse to changes in different components of bio-diversity, or that change in either direction is inherently‘‘good’’ or ‘‘bad.’’

Sustainability refers to the capacity for a given eco-system service to persist at a given level for a longperiod of time (Lubchenco et al. 1991, Valiela et al.2000). While sustainability has been discussed widely,

very few experiments have addressed it directly, in partbecause of the complexities involved. Because manyecosystem properties fluctuate naturally over time, thedifficult task is to determine the bounds of natural fluc-tuations to better understand whether human-inducedfluctuations are outside these natural ranges of vari-ability and therefore present a new threat to sustain-ability of ecosystem services (Chapin et al. 1996c).

C. Effects of diversity in the context of otherecosystem factors

Many factors influence the magnitude and stabilityof ecosystem properties, including climate, geography,and soil or sediment type. These abiotic controls in-teract with functional traits of organisms to controlecosystem properties (Fig. 1; Chapin et al. 1997, 2000,2002, Lavorel and Garnier 2002). The last half-centuryof ecosystem ecology research has yielded largeamounts of information about how organismal traitsinfluence ecosystem properties in both terrestrial andaquatic ecosystems, and about trade-offs and linkagesof these traits in individual organisms (plant effects onsoil properties, Muller 1884, Jenny 1941, 1980, VanCleve et al. 1991; species’ effects on ecosystem prop-erties, Chapin et al. 1986, 2002, Vitousek 1986, 1990,Hobbie 1992, Jones and Lawton 1995, Smith et al.1997; food webs, Carpenter et al. 1987, Carpenter andKitchell 1993, de Ruiter et al. 1994, 1995, Elser et al.1996, Schindler et al. 1997; trade-offs in plant traits,Grime 1979, Chapin 1980, Berendse et al. 1987, Grimeet al. 1988, Tilman 1988, Aerts et al. 1990, Berendseand Elberse 1990, Chapin et al. 1993, Dıaz et al. 1999,to name just a few). Ecosystem ecologists have tradi-tionally focused on the functional traits of the mostdominant organisms (those that are most abundant orhave the greatest biomass within each trophic level)because they are the most obvious biotic factors reg-ulating ecosystem properties (Grime 1998) (Box 1). Ofcourse, certain species, although relatively rare or oflow total biomass, can also have large effects (see re-view of keystone species in Power et al. 1996). In thecontext of species extinctions and invasions, under-standing the effects of diversity adds another dimensionto controls over ecosystem properties in diverse naturalecosystems (Kennedy et al. 2002). That is, under whatcircumstances do the traits of more than just one dom-inant species have a large influence on properties?When might species interactions be important? Howmany species are involved in particular community orecosystem functions? Which species play significantroles and which do not? These questions have beenstudied to some extent in agriculture and agroforestryin the context of intercropping, although the levels ofdiversity examined are usually low relative to those in


FIG. 1. Feedbacks between human activities, global changes, and biotic and abiotic controls on ecosystem properties. Anumber of human activities are now sufficiently widespread that their ecological effects have reached global proportions.These ecological effects alter both the biotic community and abiotic interactive controls on ecosystem properties. Some ofthe abiotic controls could also be considered ecosystem properties of interest. ‘‘Modulators’’ are abiotic conditions thatinfluence process rates (e.g., temperature and pH) but are not directly consumed in the process, in contrast to resources(Chapin et al. 2002). Various of aspects of the biotic community influence the range and proportion of species traits. Thesetraits can further alter the abiotic controls, directly affect ecosystem properties, or directly affect ecosystem goods andservices. Changes in ecosystem properties can feed back to further alter the biotic community either directly or via furtheralterations in abiotic controls (dotted lines). Feedbacks from altered goods and services can lead to modification of humanactivities, as evidenced in a variety of responses to environmental problems. A critical question is whether the rates andmagnitudes of these human changes will be sufficient to offset some of the original adverse ecological effects. This figureis modified from Chapin et al. (2000).

natural ecosystems (Trenbath 1974, 1999, Vandermeer1990, Swift and Anderson 1993).

Changes in biota can have greater effects on eco-system properties than changes in abiotic conditions(e.g., Van Cleve et al. 1991, Chapin et al. 2000). Im-pacts of invasions, for example, clearly demonstratethat a single species or functional group can stronglyinfluence ecosystem properties (e.g., Mooney andDrake 1986, Vitousek 1986, Griffin et al. 1989, Vitou-sek and Walker 1989, D’Antonio and Vitousek 1992,Alban and Berry 1994, Gordon 1998, Levine et al.2003). On the other hand, cross-system comparisonssuggest that abiotic conditions, disturbance regime, andfunctional traits of dominant plant species have a great-er effect on many ecosystem properties than does plantspecies richness (e.g., Wardle et al. 1997b, 2003, Lo-reau 1998a, Enquist and Niklas 2001). Modificationsof species diversity and composition result from a va-riety of environmental changes, including changes inland use, nutrient availability and cycling, atmosphericcomposition, climate, the introduction of exotic spe-

cies, and overexploitation by humans (Fig. 1). Differenttypes of environmental change are hypothesized to leadto different patterns of biodiversity modification fordifferent types of species and ecosystems (Sala et al.2000). An important goal of future research is to im-prove our understanding of the relative importance ofthe changes in different abiotic and biotic controls overspecific ecosystem properties in different ecosystems.Success in answering these questions requires a closercoupling of recent theoretical and experimental ap-proaches with the substantial information availablefrom physiological, population, community, and eco-system ecology on which sets of traits influence speciesdistributions, species interactions, and particular as-pects of ecosystem functioning (Box 1).

II. EFFECTS OF DIVERSITY ON ECOSYSTEM

PROPERTIES

A. Magnitudes of ecosystem properties

1. Theory and hypotheses.—Magnitudes of ecosys-tem processes or sizes of pools could respond to chang-


FIG. 2. Theoretical examples of how changing speciesdiversity could affect ecosystem properties. Lines show av-erage response, and points show individual treatments. (A)Selection effect for a dominant species: average ecosystemproperties increase with increasing species richness, but max-imal response is also achievable with particular combinationseven at low diversity. The increase in average response resultsfrom the greater probability of including the most effectivespecies as species richness increases. The figure illustratesresults for productivity as change in aboveground biomass.(B) Complementarity and/or positive interactions among spe-cies, illustrated for plant cover as an index of abovegroundprimary productivity in a system with all new abovegroundgrowth each year. Once there is at least one of each differenttype of species or functional type, effects of increasing spe-cies richness on ecosystem properties should begin to satu-rate; adding more species at that point would have progres-sively less effect on process rates (Tilman et al. 1997b, Loreau2000). Where the relationship saturates depends on the degreeof niche overlap among species (Petchey 2000, Schwartz etal. 2000). The figures are from Tilman (1997b).

es in species or functional diversity in several ways.The patterns depend on the degree of dominance of thespecies lost or gained, the strength of their interactionswith other species, the order in which species are lost,the functional traits of both the species lost and thoseremaining, and the relative amount of biotic and abioticcontrol over process rates (Vitousek and Hooper 1993,Lawton 1994, Naeem et al. 1995, Sala et al. 1996,Naeem 1998). Indeed, more than 50 potential responsepatterns have been proposed (Loreau 1998a, Naeem2002b). Here we focus on the most common ones andhighlight several key points.

(a) Diversity might have no effect: changing relativeabundance or species richness might not change pro-cess rates or pool sizes.—Lack of response could occurfor several reasons, such as primary control by abioticfactors, dominance of ecosystem effects by a singlespecies that was not removed, or strong overlap of re-source use by different species (Vitousek and Hooper1993, Cardinale et al. 2000, Petchey 2000, Fridley2001).

(b) Increases in ecosystem functioning with increas-ing diversity could arise from two primary mecha-nisms.—

(i) First, only one or a few species might have alarge effect on any given ecosystem property. Increas-ing species richness increases the likelihood that thosekey species would be present (Aarssen 1997, Huston1997, Tilman et al. 1997b, Loreau 2000). This is knownas the sampling effect or the selection probability ef-fect. As originally formulated, the sampling effect hy-pothesis assumes that competitive success and highproductivity are positively associated at the species lev-el (Fig. 2A; Hector et al. 2000b, Troumbis et al. 2000,Tilman 2001). Predicting the species that will have thegreatest influence on properties in complex mixtures isnot always straightforward, however. In some environ-ments, competitive success may be more stronglylinked to storage allocation, interference competition,or other strategies that do not maximize growth rates(e.g., Grime 1979, Haggar and Ewel 1995, Grime 2001,Hooper and Dukes 2004), in which case sampling ef-fects could actually lead to lower average productivity(Hector et al. 2000b, Troumbis et al. 2000, Tilman2001). For other properties, relatively rare speciescould have dominant effects on ecosystem functioning,despite having low total productivity, biomass, or abun-dance (e.g., resistance to invasions; Lyons andSchwartz 2001). Generally, we need to understandwhich traits determine competitive success and poten-tial for dominance over ecosystem properties, partic-ularly for processes other than biomass production.

(ii) Second, species or functional richness could in-crease ecosystem properties through positive interac-tions among species. Complementarity and facilitation

are the two primary mechanisms leading to the phe-nomenon of overyielding, in which plant production inmixtures exceeds expectations based on monocultureyields (Trenbath 1974, Harper 1977, Ewel 1986, Van-dermeer 1989, Loreau 1998b, but see also Petchey2003). Complementarity results from reduced interspe-


cific competition through niche partitioning. If speciesuse different resources, or the same resources but atdifferent times or different points in space, more of thetotal available resources are expected to be used by thecommunity (Trenbath 1974, Harper 1977, Ewel 1986,Vandermeer 1989). If those resources limit growth, thenincreasing functional richness should lead to greater totalproductivity and decreased loss of resources from theecosystem. Facilitative interactions among species couldalso lead to increases in ecosystem pools or process ratesas species or functional richness increase. Such facili-tation could occur if certain species alleviate harsh en-vironmental conditions or provide a critical resource forother species (Fowler 1986, Bertness and Callaway1994, Chapin et al. 1994, Berkowitz et al. 1995, Mulderet al. 2001, Bruno et al. 2003).

(c) A saturating response of ecosystem properties toincreasing species richness is the most commonly hy-pothesized pattern.—Complementarity, facilitation,and sampling effects for high productivity (or otherproperties) are all expected to show a similar saturatingaverage response as diversity increases (Fig. 2). Dis-tinguishing among these different hypotheses requirescomparisons of individual species’ performances inmonocultures and mixtures (Trenbath 1974, Tilman etal. 1997b, Hector 1998, Hooper 1998, Loreau 1998b,Mikola and Setala 1998b, Norberg 2000, Loreau andHector 2001, Drake 2003) grown close to natural den-sities to avoid yield dependence on density at low den-sity (Connolly 1986, Cousens and O’Neill 1993) anddifficulties with establishment at high densities (Harper1977). Loss of complementarity or facilitation will bemost likely to affect ecosystem properties after speciesloss has resulted in highly impoverished communities.At the same time, variability in ecosystem response tospecies loss may be expected to increase as commu-nities become more biotically impoverished because of‘‘idiosyncratic’’ effects (sensu Lawton 1994, Naeem etal. 1995) determined by the traits of the particular spe-cies going extinct or remaining in the community(Petchey 2000; see Section II.B.).

(d) Complementarity and selection or sampling ef-fects are not necessarily mutually exclusive.—Therecan be a continuum of diversity effects, ranging fromthe probability of sampling one dominant species tothe probability of selecting several complementaryspecies (Huston et al. 2000, Loreau 2000). More di-verse communities are more likely to include a dom-inant species or a particular combination of speciesthat are complementary. Furthermore, differences inresource allocation, resource use efficiency, and theamount of difference in functional traits among spe-cies could modify both complementary and samplingeffects (Haggar and Ewel 1995, Huston 1997, Tilman

et al. 1997b, Nijs and Impens 2000, Nijs and Roy2000).

(e) Ecologists disagree over whether sampling ef-fects are relevant to natural ecosystems.—Some ecol-ogists argue that they are artifacts of certain experi-mental designs because of their dependence upon anassumption that communities are random assemblagesof species from the total species pool (Huston 1997,Wardle 1999), while communities are arguably not ran-dom assemblages of species (Connell and Slatyer 1977,Weiher and Keddy 1999b). Others assert that they aresimply an alternative mechanism by which speciesrichness might influence ecosystem properties in nat-ural communities, pointing out that there are many sto-chastic factors that can influence community compo-sition (Tilman et al. 1997b, Loreau 2000, Mouquet etal. 2002). Resolving disagreements about the relevanceof sampling effects to natural systems will require abetter understanding of the links between ecosystemproperties and the interactions between deterministic(competition, trait/environment linkages) and stochas-tic (disturbance and colonization) processes that de-termine community composition.

(f) Adding multiple trophic levels is expected to leadto more complex responses in ecosystem propertiesthan in single-trophic-level models.—Most theoreti-cal research has focused on within-trophic group di-versity, such as plant diversity. Relatively few theo-retical studies examine effects of species richness onecosystem properties in multi-trophic systems (John-son 2000, Loreau 2001, Holt and Loreau 2002, The-bault and Loreau 2003). These studies suggest vari-able responses of primary and secondary productivityto changing species richness in multiple trophic lev-els, depending on a variety of factors, such as thedegree to which the system is closed to immigration,emigration, and allochthonous inputs, the degree oftop-down vs. bottom-up control, food web connectiv-ity, and the trophic level and functional characteristicsof the species gained or lost.

2. Experiments and observation.—Much of the ex-perimental work on the effects of plant diversity onecosystem properties has focused on primary produc-tivity and ecosystem nutrient retention, although agrowing number of studies have considered decom-position and nutrient dynamics as well. Intercroppingand agroforestry research is highly relevant to under-standing diversity effects on ecosystem properties (e.g.,Trenbath 1974, Harper 1977, Ewel 1986, Vandermeer1990, Loreau 1998b, Fridley 2001, Hector et al. 2002),although most such studies deal with only two to threespecies, rather than the greater diversity characteristicof natural ecosystems (Swift and Anderson 1993). Re-cent ecological experiments on the response of pro-ductivity to changing species richness in relatively di-


verse communities have focused on broader theoreticalquestions, rather than specific management goals, andsought to investigate patterns of ecosystem responsethat might occur at higher levels of diversity. Suchquestions include: What is the general shape of theresponse of productivity and other properties to in-creasing numbers of species, ranging from one speciesup to the levels of diversity characteristic of naturalcommunities? If the response saturates, at what levelof richness does this occur? What are the relative rolesof functional diversity and species diversity in affectingthat response? Many studies explicitly vary plant spe-cies richness in experimental communities in grass-lands because they are easy ecosystems to manipulateand aboveground net primary productivity is relativelyeasy to approximate because all aboveground biomassis generally accrued during a single year. Still, thesemeasurements may underestimate productivity if theydo not take into account intra-annual turnover (Scur-lock et al. 2002). Recently, evidence for properties oth-er than production and from ecosystems other thangrasslands has begun to accumulate as well, resultingin the following generalizations:

(a) Differences in species composition exert a strongeffect on productivity and other ecosystem proper-ties.—Ecosystem response to extinction or invasion inthe real world will be determined at least as much bywhich species and functional traits are lost and remainbehind as by how many species are lost. As statedabove (Section I.C.), research in ecosystem ecologyover the past half century has demonstrated that or-ganismal functional traits are one of the key controlson ecosystem properties. Recent studies on the effectsof diversity on ecosystem functioning in both terrestrialand aquatic ecosystems support those findings: Mostobserve large variability in ecosystem properties withinlevels of species or functional richness that can be at-tributed at least in part to differences in species orfunctional composition (Fig. 3; Naeem et al. 1995, Til-man et al. 1996, 1997a, Haggar and Ewel 1997, Hooperand Vitousek 1997, Hooper 1998, Symstad et al. 1998,Hector et al. 1999, Norberg 1999, Wardle et al. 1999,Spehn et al. 2000, Van der Putten et al. 2000, Leps etal. 2001, Hector 2002). These experiments suggest that,as predictors of ecosystem properties, community com-position (knowing which species or functional typesare present) is at least as important as species or func-tional richness alone (knowing how many species orfunctional types are present).

Soil processes in particular appear to be primarilyinfluenced by the functional characteristics of dominantspecies rather than by the number of species present(but see Zak et al. 2003). Decomposition, soil organicmatter dynamics, nutrient uptake by soil micro-organ-isms, and nutrient retention, for example, are more

strongly influenced by differences in functional traits(e.g., leaf chemistry, phenology) of the dominant plantspecies than by the diversity of plant species (Hooperand Vitousek 1997, 1998, Wardle et al. 1997a, b, 1999,Bardgett and Shine 1999, Hector et al. 2000a, Korthalset al. 2001). Less is known about how the diversity ofsoil organisms affects rates of decomposition and nu-trient cycling (Balser et al. 2002, Mikola et al. 2002).Composition and diversity of mycorrhizal fungi influ-ence plant community composition and productivity(van der Heijden et al. 1998, 1999, but see Wardle1999), as well as productivity of individual plants, buteffects can be positive, negative, or neutral dependingon soil fertility and the plant species involved (Jonssonet al. 2001). Litter decomposition rates can depend onthe composition of the soil faunal community, whichin turn is influenced by the plant species present (Chap-man et al. 1988, Blair et al. 1990, Williams 1994, butsee also Andren et al. 1995). Experimental studiesbased on synthesized soil food webs point to food webcomposition, rather than the diversity of organismswithin trophic levels, in driving decomposition prop-erties (Mikola and Setala 1998a) and plant productivity(Laakso and Setala 1999).

(b) Patterns of response to experimental manipula-tions of species richness vary for different processes,different ecosystems, and even different compartmentswithin ecosystems.—In some experiments with herba-ceous plants, average plant productivity increases, andlevels of available soil nutrients often decrease, withincreasing plant species or functional richness, at leastwithin the range of species richness tested and over therelatively short duration of many experiments (Fig. 3;Tilman et al. 1996, 1997a, 2001, 2002, Hector et al.1999, Loreau and Hector 2001, Niklaus et al. 2001a,Fridley 2003). In these experiments, the responses tochanging diversity are strongest at low levels of speciesrichness and generally saturate at 5–10 species (but seeSection II.B., below, for more on levels of saturation).However, increases in process rates with increasing spe-cies richness do not always occur. In some experimentswith longer-lived perennials, ecosystem responses (NPP,nutrient retention, nutrient use efficiency) are maximizedwith only one or two species (e.g., Ewel et al. 1991,Haggar and Ewel 1997, Hiremath and Ewel 2001; Fig.3). Idiosyncratic patterns sometimes result from strongeffects of species composition, in which the functionaltraits of particular species overwhelm responses to spe-cies richness (Hooper and Vitousek 1997, Symstad etal. 1998, Kenkel et al. 2000, Troumbis et al. 2000,Mulder et al. 2001). These patterns, seen under exper-imental conditions, may or may not reflect actual pat-terns seen for a particular ecosystem under a particularscenario of species loss or invasion, which will dependnot only on the functional effect traits of the species


FIG. 3. Variation in effects of plant species richness and composition on plant productivity. (A) Experiments in the tropics.Treatments ran for five years and included four monocultures (two rotations [1st and 2nd] of maize [Zea mays], one rotationof cassava [Manihot esculenta], and one rotation of a tree, Cordia alliodora); a diverse (.100 plant species) natural successionfollowing clearing and burning of original vegetation; a species-enriched (;120 species) version of natural succession; andan imitation of succession that mimicked the plant life forms in the natural succession treatment, but with different species.Monocultures were timed to coincide with growth phases of natural succession: maize during the initial herbaceous stage,cassava during the shrub-dominated stage, and C. alliodora during the tree-dominated stage. Note that the maize monoculturehad both the highest and lowest overall productivity, and that the productivity of the successional vegetation was not increasedby further increases in species richness. This figure is modified from Ewel (1999). (B) The pan-European BIODEPTHexperiment. At several sites, plant productivity increased with increasing species richness, although the pattern of responsevaried in individual location analyses. Five of the sites had either non-saturating or saturating patterns (on a linear scale).At two sites significant differences across different levels of diversity (ANOVA) provided a better model than a linearregression. One site (Greece, dotted line) showed no significant relationship between aboveground plant productivity andspecies richness. Even where there are strong trends in the diversity effect, there is also variation within levels of richnessresulting in part from differences in composition. Points are individual plot biomass values, and lines are regression curvesor join diversity level means (squares for Ireland and Silwood). The figure is after Hector et al. (1999).


involved, but also on the traits that determine how spe-cies respond to changes in environmental conditions(i.e., both effect and response traits, Box 1 and SectionII.B.2; Symstad and Tilman 2001, Lavorel and Garnier2002). Understanding the causes of variability in re-sponse patterns for different ecosystem types and dif-ferent environmental conditions remains an importantquestion.

(c) Both sampling effects and positive species inter-actions have been observed in experiments, and mul-tiple mechanisms can operate simultaneously or se-quentially.—Resolving the mechanisms by which ex-perimental manipulation of species richness leads toincreased productivity or other processes has led tosubstantial debate (see Section II.A.1. Theory and hy-potheses, above; Aarssen 1997, Garnier et al. 1997,Huston 1997, van der Heijden et al. 1999, Wardle 1999,Hector et al. 2000b, Huston et al. 2000). Many exper-iments were designed to test general patterns, ratherthan to test mechanisms for those patterns. Those thatdo test explicitly for mechanism clearly indicate thatalternatives are often not mutually exclusive. Both pos-itive interactions among species (complementarity and/or facilitation) and selection for highly productive spe-cies occurred in synthetic grassland communities inEurope (Loreau and Hector 2001) and Minnesota (Til-man et al. 1996, 2001, Reich et al. 2001).17 Positiveinteractions involving at least two species are occur-ring, but whether this results from facilitation or com-plementarity and how many species are involved isunclear (Huston and McBride 2002, Tilman et al. 2002,Wardle 2002). Evenness of the plant community alsocould lead to increased productivity with increasingspecies richness (e.g., Nijs and Roy 2000, Schwartz etal. 2000, Wilsey and Potvin 2000, Polley et al. 2003).Effects of plant diversity on soil nutrients can be me-diated simultaneously by direct plant uptake and byeffects of plants on soil microbial dynamics (Hooperand Vitousek 1997, 1998, Niklaus et al. 2001a).

Several questions remain unresolved. For example,what functional traits of species lead to dominance andhow do traits for dominance overlap with functionaleffect traits (Weiher and Keddy 1999b, Suding et al.2003)? Several recent experiments have shown that the

17 Note that the use of the term ‘‘selection’’ in the AdditivePartitioning Equation (APE; Loreau and Hector 2001) is dif-ferent from the ‘‘sampling effect’’ or ‘‘selection probabilityeffect’’ (Huston 1997, Tilman et al. 1997b). The ‘‘selectioneffect’’ of the APE refers to the tendency for species inter-actions in mixtures to ‘‘select for’’ or favor species with par-ticular traits (e.g., high productivity in monoculture), whereassampling effects refer to the higher probability of includingsuch species in randomly selected mixtures as the speciesrichness of experimental treatments increases. Both of theseaspects must hold for sampling effects to be the primarydriver of ecosystem properties.

species with the greatest productivity in monocultureis not necessarily the species that dominates productionin mixtures (Hooper and Vitousek 1997, Troumbis etal. 2000, Engelhardt and Ritchie 2001, Spaekova andLeps 2001, Hector et al. 2002, Hooper and Dukes2004), contrary to some early formulations of the sam-pling effect hypothesis (Huston 1997, Tilman et al.1997b).

To further understand diversity effects on ecosystemproperties, future experiments need to include explicitexperimental controls (e.g., growing all species inmonoculture as well as in mixture, Hector 1998, Hoop-er 1998, Loreau 1998b, Engelhardt and Ritchie 2002,Fridley 2003, Hooper and Dukes 2004; or having ma-trix species alone at different densities, Haggar andEwel 1997), or at the minimum, statistical controls(e.g., measurements of potential controlling variables)to help differentiate among mechanisms (Huston andMcBride 2002, Schmid et al. 2002a). Optimally, grow-ing all possible polycultures, as well as monocultures,would help distinguish sampling effects for small num-bers of species, but this approach may not be experi-mentally tractable.

(d) The strength of positive interactions varies withboth the functional characteristics of the species in-volved and the environmental context.—Extensive re-search over many decades in intercropping and agro-forestry shows that the degree of complementarity orfacilitation among crop or forestry species varies great-ly (e.g., Vandermeer 1989, 1990, Ong and Black 1994,Haggar and Ewel 1997, Ong and Huxley 1997). Similarvariation in the strength of positive interactions occursin ecological experiments, such as those investigatingcompetition (Harper 1977, Berendse 1982, 1983, Baz-zaz 1987), and more recent experiments assessing com-plementarity and facilitation among terrestrial andaquatic plants (Hooper 1998, Dukes 2001b, Engelhardtand Ritchie 2002, Schmid et al. 2002b, Fridley 2003,Polley et al. 2003, van Ruijven and Berendse 2003,Hooper and Dukes 2004), and aquatic animals (Norberg2000, Emmerson et al. 2001).

Complementarity and/or facilitation are usuallygreatest when species differ greatly in functional traits,whether in timing (Steiner 1982, Chesson et al. 2002,but see Stevens and Carson 2001), spatial distribution(Schenk and Jackson 2002), or type of resource demand(e.g., McKane et al. 2002). One of the most importantforms of facilitation among plants occurs when at leastone species has the ability to form a symbiotic asso-ciation with nitrogen-fixing bacteria (Trenbath 1974,Cannell et al. 1992). Interactions between legumes andnon-legumes are clearly one of the major functionalmechanisms for the results of many grassland biodi-versity experiments (e.g., Tilman et al. 1997a, 2002,Hector et al. 1999, Mulder et al. 2002, Spehn et al.


2002). However, they are not necessarily the wholestory. Effects of additional species can be detected insome of these studies (Loreau and Hector 2001, Tilmanet al. 2001, 2002) and overyielding has been found inmany mixtures omitting legumes (e.g., Trenbath 1974,Haggar and Ewel 1997, Jolliffe 1997, van Ruijven andBerendse 2003, Hooper and Dukes 2004).

Environmental context, both abiotic and biotic, canadd variability to the strength of positive interactions(Cardinale et al. 2000, Emmerson et al. 2001, Fridley2001, Hooper and Dukes 2004). In intercropping stud-ies, much effort goes into determining the appropriateconditions (e.g., spacing of individuals, timing of plant-ings, soil conditions) to maximize total yields. In nat-ural systems, facilitation is most common in unpro-ductive or stressful environments (Bertness and Cal-laway 1994, Callaway et al. 2002, Bruno et al. 2003).On the other hand, increasing resource availability mayallow for stronger complementarity. Positive short-term effects of species richness on aboveground pro-ductivity are often greater with higher resource avail-ability, such as CO2 or fertilizer enrichment (Stockeret al. 1999, Niklaus et al. 2001b, Reich et al. 2001,Fridley 2002, 2003, He et al. 2002), although evidencesuggests both complementarity and sampling effects asthe underlying mechanisms in different experiments.Such results need to be reconciled with the well-knownphenomenon of decreasing plant diversity with increas-ing fertilization (e.g., Grime 1973a, 1979, Tilman1987). For example, how do predictions for positiveinteractions relate to predictions from the humpbackedmodel of species diversity (see Sections I.C. and II.C.)?The influence of environmental variation and differ-ences in species’ functional traits on complementarityand facilitation in complex natural and seminaturalcommunities deserves more empirical study.

(e) Higher species richness within sites tends to de-crease invasion by exotic species, though cross-sitecomparisons often show positive correlations betweenrichness and invasibility.—At the landscape-scale, var-iability in factors such as soil fertility, propagule input,and disturbance regimes tend to outweigh effects ofspecies richness on invader success, often leading topositive correlations between invader success and spe-cies richness when making comparisons across differ-ent sites (Planty-Tabacchi et al. 1996, Levine andD’Antonio 1999, Stohlgren et al. 1999, Levine 2000),although counterexamples exist (Gido and Brown1999, Sax and Brown 2000). However, when makingcomparisons under common conditions, increasing spe-cies richness generally decreases the success of inva-sives (McGrady-Steed et al. 1997, Tilman 1997a, 1999,Knops et al. 1999, Stachowicz et al. 1999, Levine 2000,Naeem et al. 2000b, Prieur-Richard and Lavorel 2000,Symstad 2000, Dukes 2001a, Hector et al. 2001a, Ly-

ons and Schwartz 2001, Kennedy et al. 2002, Fargioneet al. 2003). A decrease in invasibility with increasingspecies richness within sites could occur by a varietyof mechanisms, such as a greater probability of in-cluding species with traits similar to potential invaders,by more species utilizing a greater proportion of thepotentially available resources (Elton 1958, Tilman1999), a greater probability of including strongly com-petitive species (Wardle 2001a), or the greater likeli-hood of including biotic controls of a prospective in-vader. Conversely, increasing species richness can in-crease invasibility within sites if these additions resultin increased resource availability, as in the case of ni-trogen-fixers (Prieur-Richard et al. 2002a), or increasedopportunities for recruitment through disturbance (e.g.,D’Antonio 2000). Integrating results from field surveyswith results from within-site experimental manipula-tions and mathematical models is important for boththeoretical understanding and for broad-scale manage-ment of exotic species’ invasions (Levine andD’Antonio 1999, Levine 2000, Shea and Chesson2002).

(f) Varying diversity and composition of hetero-trophs can lead to more idiosyncratic behavior thanvarying diversity of primary producers alone.—Asmultitrophic diversity increases, average process ratescould increase, decrease, stay the same, or follow morecomplex nonlinear patterns (e.g., Carpenter and Kitch-ell 1993, Schindler et al. 1997, Klironomos et al. 2000,Cardinale et al. 2002, Mikola et al. 2002, Paine 2002,Raffaelli et al. 2002; see also Section II.A.1(f), above).Such complex patterns (e.g., Thebault and Loreau2003) might explain why experimental results obtainedwith a small number of diversity levels appear some-what variable. Many aquatic and terrestrial experimentshave manipulated the abundance of one or a few con-sumer species (citations in previous sentences). Agrowing number of experiments have specifically ma-nipulated diversity of more than one trophic level, al-though experimental difficulties in doing so restrictmany of these experiments to micro- or mesocosms(e.g., Naeem et al. 1994, 2000a, McGrady-Steed et al.1997, Mikola and Setala 1998a, Laakso and Setala1999, Mulder et al. 1999, Petchey et al. 1999, Wardleet al. 2000a, Downing and Leibold 2002, plus abovereferences).

The major point that emerges is that the functionalcharacteristics of single species, whether native or not,can have a large impact on both community structureand ecosystem functioning. Changes in compositionand diversity at one trophic level can influence diver-sity either positively or negatively in other trophic lev-els by a variety of mechanisms (Hunter and Price 1992,Strong 1992, Wardle et al. 1999, Duffy and Hay 2000,Hooper et al. 2000, Klironomos et al. 2000, Norberg


2000, Stephan et al. 2000). Subtle differences in speciesinteractions and environmental conditions can make theresulting effects on ecosystem properties difficult topredict (Berlow 1999, Wolters et al. 2000, Duffy et al.2001, Schmid et al. 2002b). However, understandingthe functional relationships among species within andacross trophic levels helps to explain some of the ap-parently idiosyncratic ecosystem behavior that results(de Ruiter et al. 1994, 1995, Hulot et al. 2000, Bradfordet al. 2002). Greater experimental efforts at understand-ing multitrophic changes in diversity constitute a clearneed for future research.

B. Variability in ecosystem properties

1. Theory and hypotheses.—Ecologists hypothesizethat ecosystem properties should be more stable in re-sponse to environmental fluctuations as diversity in-creases. Studies of the relationship between diversityand stability have a long tradition in ecology (Mac-Arthur 1955, May 1974, Pimm 1984, McCann 2000),but findings have sometimes been clouded by incon-sistent terminology. First, the distinction must be drawnbetween the stability of community composition andthe stability of ecosystem process rates. In the formercase, changing community composition is consideredinstability (May 1974); in the latter case, changingcommunity composition is one mechanism that canhelp promote stability of ecosystem properties (Mc-Naughton 1977, Tilman 1996, 1999, Lehman and Til-man 2000). In addition, ‘‘stability’’ in biotic commu-nities is an umbrella term that refers to a large numberof potential phenomena, including, but not limited to,resistance to disturbance, resilience to disturbance,temporal variability in response to fluctuating abioticconditions, and spatial variability in response to dif-ferences in either abiotic conditions or the biotic com-munity (May 1974, Pimm 1984, Holling 1986, Mc-Naughton 1993, Peterson et al. 1998, Chesson 2000,Lehman and Tilman 2000, Cottingham et al. 2001,Chesson et al. 2002, Loreau et al. 2002a). Most the-oretical work has focused on temporal variability, al-though some of the same principles may apply to othertypes of stability. Exploring the effects of species rich-ness and composition on other dimensions of stabilityis a clear need for future research.

Theory about the relationship between species rich-ness and stability of ecosystem processes has been de-veloped in several forms, both via simple ecologicalreasoning and via mathematical models. Consensus onseveral points emerges from these different approaches.

(a) A diversity of species with different sensitivitiesto a suite of environmental conditions should lead togreater stability of ecosystem properties.—In thissense, redundancy of functional effect traits and di-versity of functional response traits (see Box 1) act as

insurance in carrying out ecological processes (Mac-Arthur 1955, Elton 1958, Chapin and Shaver 1985,Walker 1992, Lawton and Brown 1993, Naeem 1998,Petchey et al. 1999, Trenbath 1999, Walker et al. 1999,Yachi and Loreau 1999, Hooper et al. 2002). If anecosystem is subject to a variety of natural and human-caused environmental stresses or disturbances, thenhaving a diversity of species that encompass a varietyof functional response types ought to reduce the like-lihood of loss of all species capable of performing par-ticular ecological processes, as long as response traitsare not the same as or closely linked to effect traits(Chapin et al. 1996a, Lavorel and Garnier 2002). Thisdiversity of different functional response types alsoleads to asynchrony in species’ demographic responsesto environmental changes. Asynchrony results in com-pensation among species: As some species do worse,others do better because of different environmental tol-erances or competitive release. In such cases, unstableindividual populations stabilize properties of the eco-system as a whole (McNaughton 1977, Tilman 1996,1999, Hughes and Roughgarden 1998, Ives et al. 1999,Landsberg 1999, Walker et al. 1999, Lehman and Til-man 2000, Ernest and Brown 2001a). By similar rea-soning, processes that are carried out by a relativelysmall number of species are hypothesized to be mostsensitive to changes in diversity (Hooper et al. 1995),and loss of regional species richness is hypothesizedto compromise recruitment and regeneration of poten-tially dominant species under changing environmentalconditions (Grime 1998).

Several mathematical models generally agree withthe hypotheses just described (see McCann 2000, Cot-tingham et al. 2001, Loreau et al. 2002a, for reviews).If species abundances are negatively correlated or varyrandomly and independently from one another, thenoverall ecosystem properties are likely to vary less inmore diverse communities than in species-poor com-munities (Fig. 4; Doak et al. 1998, Tilman et al. 1998).This statistical averaging is similar to diversified stockportfolios: The more companies in which one invests,the lower the risk of losing all of one’s savings shouldone company collapse. The strength of the modeledeffects of asynchrony depends on many parameters,including the degree of correlation among differentspecies’ responses (Doak et al. 1998, Tilman et al.1998, Tilman 1999, Yachi and Loreau 1999, Lehmanand Tilman 2000, Chesson et al. 2002), the evennessof distribution among species’ abundances (Doak et al.1998), and the extent to which the variability in abun-dance scales with the mean (Tilman 1999, Yachi andLoreau 1999, Cottingham et al. 2001).

(b) The numbers of species or genotypes necessaryto maintain ecosystem properties increases with in-creasing spatial and temporal scales.—It follows from


FIG. 4. Simulated stability of individual populations and a resulting community property (summed abundances of indi-vidual species) illustrating the portfolio effect of Doak et al. (1998). The model is from Tilman (1999); the figure is fromCottingham et al. (2001). The decrease in aggregate variability with increasing numbers of species results from the randomfluctuations of the individual species. Underlying assumptions that contribute to the degree of dampening include equalabundance of all species and no correlation (r 5 0) among species’ temporal dynamics.

point 1, above, that, while magnitudes of ecosystemproperties may saturate at relatively low levels of spe-cies richness in small-scale, short-term experiments,more genetic diversity, either in terms of different spe-cies or genetic diversity within species, is necessary asa greater variety of biotic and abiotic conditions areencountered (Field 1995, Pacala and Deutschman 1995,Casperson and Pacala 2001, Chesson et al. 2002). Thiscould have a variety of implications for the sustain-ability of ecosystem services in the long term (see Sec-tion III, below; Ewel 1986).

(c) The underlying assumptions of the mathematicalmodels need further investigation and more experi-mental confirmation.—These assumptions include thedegree of negative covariance, the relative abundancesof species, the measures of stability used, and theamount of overyielding built into the models (Cottingh-am et al. 2001, Chesson et al. 2002). To that end, newor different models that encode these same assumptionsdo not necessarily lend more support to the diversity/stability hypothesis; they are simply different mathe-matical configurations of the same thing. For example,several models have negative covariance, equal speciesabundances, or overyielding built in, either implicitly

or explicitly (e.g., Lehman et al. 1975, Tilman 1999,Lehman and Tilman 2000). Increasing productivitywith increasing species richness via overyielding leadsto greater stability if the coefficient of variation (CV)or its inverse, S (Tilman 1999) is used as the measureof stability, because of a higher mean productivity, notbecause of lower variance (Lehman and Tilman 2000).The strength of stabilization is likely to be maximal insuch cases (Doak et al. 1998, but see also Yachi andLoreau 1999). Further exploration of the parameterspace for all these variables is necessary before suchmodels can be considered more proof that diversitystabilizes ecosystem processes.

Similarly, use of either CV or net variance as a mea-sure of stability is well supported theoretically, butwhich measure is most relevant and the extent to whichstability might be influenced may depend on the par-ticular application and how variance scales with themean (the ‘‘z’’ scaling factor; Hughes and Roughgarden1998, Tilman 1999, Cottingham et al. 2001, but seealso Yachi and Loreau 1999). Modelers need to separateeffects of changes in the mean, variance, and covari-ance on measures of stability used (Lehman and Tilman2000). This distinction could be important for man-


agement issues where probability of loss of a functionor of maintaining a minimum level of function are con-cerned.

In short, both heuristic theory and several mathe-matical models predict that increased diversity will leadto lower variability of ecosystem properties under thoseconditions in which species respond asynchronously totemporal variation in environmental conditions. Whilethese theoretical studies provide insights about poten-tial mechanisms, they cannot tell us how importantthese mechanisms are in the real world or whether theysaturate at high or low levels of species richness. Keyassumptions about equitability of species distributionsand the degree of resource partitioning in some modelsare not necessarily realistic for many ecosystems(Schwartz et al. 2000, Cottingham et al. 2001). Furtherexploration of the parameter space as these assump-tions are relaxed would contribute greatly to our un-derstanding of the conditions under which diversitymight be expected to contribute to various aspects ofstability in real ecosystems.

2. Experiments and observations.—While theoryabout effects of species and functional diversity onstability of ecosystem properties is relatively well de-veloped, testing the predictions of this theory is moredifficult. Such studies require long-term investigationsof communities where differences in species diversityare not confounded by variation in other ecosystemproperties, such as soil fertility or disturbance regime.They require observing properties both before and afterdisturbances or strong environmental fluctuations. Andthey require many generations of the experimental or-ganisms. For example, among consumer organisms,compensation could take place by either greater percapita consumption or greater population sizes, the lat-ter of which clearly needs time to develop over multiplegenerations (Ruesink and Srivastava 2001). Because ofthese difficulties, relatively few experiments have beencarried out in the field compared to microcosm studies,in which experiments can be conducted for dozens tohundreds of generations on organisms such as microbesand small invertebrates. Microcosm experiments allowtesting of theoretical principles in relatively controlledconditions, though proof that either the theory or mi-crocosm findings apply to the real world requires morework (Naeem 2001). In addition, relatively few exper-iments, in either microcosms or the field, have beenable to completely avoid confounding the effects ofspecies richness with effects of other variables on themeasured responses. Despite these limitations, the fol-lowing consensus points emerge from experimentalstudies:

(a) In diverse communities, redundancy of functionaleffect types and compensation among species can buff-er process rates in response to changing conditions

and species losses.—Considerable evidence existsfrom field studies in a variety of ecosystems. In lakes,redundancy in species effects on ecosystem propertiesis a common feature, at least at lower trophic levels(Frost et al. 1995). For example, primary productionwas relatively constant despite changes in the numberand composition of phytoplankton species in responseto experimental acidification in a Canadian Shield lake(Schindler et al. 1986). In contrast, changes in speciesnumber and composition of higher trophic levels,which generally have lower diversity and therefore lessredundancy, often lead to major changes in both com-munity composition and productivity of lower trophiclevels in marine and freshwater ecosystems (Schindleret al. 1986, Carpenter and Kitchell 1993, Estes et al.1998, Vander Zanden et al. 1999, Lodge et al. 2000).Even in diverse communities, however, compensationmay not occur among all species in a given trophiclevel, suggesting that further refinement of functionaleffect groups beyond trophic position is necessary (Hu-lot et al. 2000, Duffy et al. 2001, Ruesink and Srivas-tava 2001).

Experiments that have tried to remove key taxonomicgroups in soil food webs have found relatively littlechange in average process rates such as soil respiration,aboveground net primary production (NPP), and netecosystem production (Ingham et al. 1985, Liiri et al.2002). The high diversity of soil organisms and therelatively low degree of specialization in detritivoresmeans that many different species can carry out similarprocesses (Bradford et al. 2002, but see also Mikola etal. 2002). Loss of redundancy within functional effectgroups and its buffering capacity for ecosystem prop-erties may not be apparent until ecosystems have beenexposed to multiple types of stresses (Griffiths et al.2000, de Ruiter et al. 2002).

In aboveground communities, changes in resourceavailability, temperature, and disturbance regime canbe buffered at the ecosystem level by shifts in speciescomposition (Chapin et al. 1996a, Walker et al. 1999).For example, changes in nutrients and temperature ledto large shifts in species composition, but relativelylittle change in total productivity in long-term exper-iments in Arctic tundra (Chapin and Shaver 1985). Var-iability in populations appeared to be at least partiallyresponsible for decreased ecosystem variability in re-sponse to water availability in Minnesota grasslands(Tilman 1996, 1999, Tilman et al. 2002; but see alsopoint (c), below). Compensation among species of de-sert rodents clearly stabilized ecosystem properties, al-though the degree of compensation and stability wasnot tested across different levels of diversity (Ernestand Brown 2001a). Studies of ecosystem recovery afterdisturbance have often found that ecosystems withmore rapid recovery (i.e., greater resilience) were those


FIG. 5. Increasing stability with increasing species rich-ness in ecological experiments. In both cases, the overallpatterns are as predicted from theory, but the underlyingmechanisms may coincide only in part (see Section II.B.2).(A) Temporal variability (coefficient of variation, CV) inaboveground plant biomass (correlated with productivity inthese Minnesota grasslands) in response to climatic variabil-ity (the figure is from Tilman [1999]). The gradient in speciesrichness results from different levels of nutrient addition, sothat the stability response may result from differences in spe-cies composition instead of, or in addition to, compensatoryresponses among species (Givnish 1994, Huston 1997). (B)Standard deviation (SD) of net ecosystem CO2 flux in a mi-crobial microcosm (the figure is from McGrady-Steed et al.[1997]). The decrease in variability with increasing diversitymay result from both decreased temporal variability and in-creased compositional similarity among replicates. See alsoMorin and McGrady-Steed (2004). Composite figure afterLoreau et al. (2001).

with a higher diversity of response types (e.g., a mixof seeders and sprouters in the case of fire; Lavorel1999).

(b) Mechanisms other than compensation can affectstability in response to changing species richness orcomposition.—Frank and McNaughton (1991) foundincreased stability of community composition at higherspecies richness in Yellowstone grasslands, though the-ory predicts the opposite (May 1974, Tilman 1999,Tilman et al. 2002). Stability of production underdrought in bryophyte communities increased with in-creasing species richness, but resulted from facilitativeinteractions rather than compensation among species(Mulder et al. 2001). Particular functional traits, suchas the degree of nutrient stress tolerance or evolution-ary history of exposure to a certain disturbance, canbe strong predictors of ecosystem and community re-sponse to disturbance, even without invoking speciesrichness or compensatory interactions (MacGillivray etal. 1995, Sankaran and McNaughton 1999, Wardle etal. 2000a). Stability to experimental drought actuallydecreased with increasing plant species richness inSwiss meadows because of positive effects of nitrogen-fixers on overall productivity, but susceptibility ofthose N-fixers to drought (Pfisterer and Schmid 2002;but see also Schmid and Pfisterer 2003, Wardle andGrime 2003). In agricultural ecosystems, genetic andspecies diversity of crops and increased diversity ofassociated insect species can reduce susceptibility ofcrops to climate variability, pests, pathogens, and in-vasion of weedy species (e.g., Trenbath 1999, Zhu etal. 2000). However, these patterns also have counter-examples. For example, natural pest control may in-crease with increasing diversity of associated plant andinsect species in some cases (Naylor and Ehrlich 1997),but in others, more diverse settings lead to greater pestpopulations, e.g., by providing key hosts of high pal-atability or that allow pests to complete a complex lifecycle (Brown and Ewel 1987, Prieur-Richard et al.2002b). Such counterexamples suggest that the rightcombinations of functional attributes, not just diversityeffects, often play a major role in determining ecosys-tem response.

(c) Several experiments that manipulate diversity inthe field and in microcosms generally support theo-retical predictions that increasing species richness in-creases stability of ecosystem properties, although mostexperiments are confounded by other variables (Fig.5).—The experimental difficulty reflects both the com-plexity of controlling a variety of potentially confound-ing variables and ecologists’ increased understandingof what those variables are. Stability of plant produc-tion, as measured by resistance and/or resilience to nu-trient additions, drought, and grazing, increased withthe Shannon-Wiener index of diversity (H9) in a variety

of successional and herbivore-dominated grasslands(McNaughton 1977, 1985, 1993). However, results ofthese early experiments may be confounded by a va-riety of factors, such as differences in species com-position and/or abiotic conditions, which also haveraised controversy in more recent experiments. For ex-ample, in Minnesota grasslands, resistance to loss ofplant productivity to drought increased with increasingplant species richness (Tilman and Downing 1994).However, because the species richness gradient in thisexperiment was caused by nutrient additions, the sta-


bility response may have resulted as much from com-position differences caused by the nutrient additions asfrom compensation among species (Leps et al. 1982,Givnish 1994, MacGillivray et al. 1995, Huston 1997,Grime et al. 2000, Pfisterer and Schmid 2002; but seeTilman et al. 1994). Experiments in microcosms andgrasslands suggest that increased species richness, ei-ther in terms of numbers of different functional groups,or numbers of species within trophic functional groups,can lead to decreased temporal variability in ecosystemproperties (McGrady-Steed et al. 1997, Naeem and Li1997, Petchey et al. 1999, Emmerson et al. 2001, Pfis-terer et al. 2004; but see also Pfisterer and Schmid2002). While species richness or H9 was statisticallysignificant in all these experiments, species composi-tion (where investigated) had at least an equally strongeffect on stability. In some experiments, effects of di-versity on temporal variability via compensation orportfolio effects were confounded with effects of com-positional similarity among replicates at higher levelsof diversity (Wardle 1998). The correlation betweencompositional similarity and species richness may re-semble situations resulting from species loss in realcommunities (Naeem 1998, Fukami et al. 2001), butdetermining mechanisms responsible for patterns ofecosystem response becomes problematic.

(d) Explicit demonstration of compensation amongspecies requires careful experimental control and can-not be taken for granted as the mechanism underlyingstability responses.—Careful consideration of the ques-tions being asked is required to assess a variety oftrade-offs in experimental design for experiments ondiversity effects on stability (and magnitudes) of eco-system properties. Important aspects of experimentaldesign include maximum species richness levels rela-tive to the size of species pool, the degree of exactreplication of composition treatments, random selec-tion of species vs. particular scenarios of communityassembly/disassembly, and types of statistical analysis(Allison 1999, Emmerson and Raffaelli 2000, Hooperet al. 2002, Huston and McBride 2002, Schmid et al.2002a).

In sum, the experimental work provides qualifiedsupport for the hypothesis that species richness canaffect stability of ecosystem properties, although theunderlying mechanisms can differ from theoretical pre-dictions and in many cases still need to be fully re-solved (Loreau et al. 2001). To this end, a closer linkingof theory and experiments would be helpful. Experi-ments and measurements in natural communitiesshould address explicit predictions and assumptions de-veloped in theoretical models. These include measuringchanges in species composition, evenness, correlationsamong population fluctuations, and values of the scal-ing factor z, as well as ecosystem properties, and com-

paring effect and response traits in intact vs. disturbedecosystems. In addition, more theoretical investigationof the measures of process stability, such as resilience,resistance, and spatial variability, in addition to tem-poral variability, would help with applicability to ex-periments. Some of the theory developed for temporalvariability may apply to other measures of ecosystemstability, but more exploration of when, where, and why(or why not) is necessary.

C. Matching experiments with observation

1. Productivity effects on diversity and vice versa.—Species composition and richness both respond to andinfluence ecosystem processes (Fig. 1). Until recently,studies of the response of diversity to variation in pro-ductivity, disturbance regime, or resource availabilityhave been much more common than studies of the re-sponse of those processes to variation in diversity.Studies of the former type assess the influence of var-iation in abiotic factors on species diversity. One pat-tern that is often observed is a unimodal relationship(the so-called ‘‘humpbacked curve’’) in which speciesrichness is greatest at intermediate levels of resourceavailability, stress, productivity, or disturbance (Grime1973b, 1979, 2001, Connell 1978, Huston 1979, 1994,Sousa 1979, Dodson et al. 2000). However, dependingon organism type, the intensity of abiotic constraints,and the geographic scale, a variety of patterns can oc-cur, including lower diversity at higher levels of pro-ductivity (Waide et al. 1999, Mittelbach et al. 2001).Furthermore, allometric scaling relationships suggestno relationship between plant diversity and total com-munity biomass across a wide variety of tree-dominatedcommunities (Enquist and Niklas 2001, Enquist et al.2002). How can these observations be reconciled withthose experiments in which primary productivity isgreater at high species richness?

Essentially, the two approaches ask different ques-tions. Macroecological patterns such as the humped-back relationship are seen in response to gradients ofabiotic factors that influence productivity, either acrosssites or within sites in response to changes in resourcesupply. Diversity then responds to the resulting abioticenvironment and represents the species functional traitsfor which that environment selects (Grime 1979, Ro-senzweig and Abramsky 1993, Ewel 1999, Mittelbachet al. 2001). On the other hand, diversity effects onproductivity are primarily observed within sites undersimilar environmental conditions (Loreau 1998a, Lo-reau et al. 2001, Schmid 2002). The way the diversity/ecosystem functioning question was originally framedimplied not cross-system comparisons, but changes indiversity within a given system in response to humanactivities. This difference explains, in part, the pre-dominantly theoretical and manipulative bent of recent


studies of effects of diversity on ecosystem properties.If the goal is to understand how changes in speciesrichness affect ecosystem properties, one needs gra-dients in diversity that are not confounded by variationin other potentially important ecosystem controls (seeSection II.C.2, below). Such diversity gradients are dif-ficult to find in natural ecosystems.

A synthesis of the local and macroscale perspectivesis needed. Anthropogenic changes that lead to speciesextinctions rarely result in similar environmental con-ditions in the real world. Community diversity andcomposition therefore are best viewed as dynamic var-iables, both responding to and affecting environmentalconditions and ecosystem properties. Extinctions ofsome species may lead to replacements or compensa-tion by others (Brown et al. 2001), and these dynamicscan themselves be strong determinants of ecosystemproperties (e.g., Ernest and Brown 2001b, Symstad andTilman 2001, Levine et al. 2003). Changes in ecosys-tem properties can then feed back to further influencespecies composition and other ecosystem properties(Fig. 1; Chapin 1980, Hobbie 1992, Chapin et al. 2000).Some evidence suggests that if experimental diversitygradients are not maintained by continuous weeding,the positive relationship between species richness andprimary productivity can decay rapidly (Pfisterer et al.2004). On the other hand, recruitment limitations thatcause time lags in colonization by new species couldslow re-establishment of a new community (Grime1998). Even if an effect of reduced diversity was tran-sient, it could still last for decades or longer for long-lived communities such as forests. A better understand-ing of the forces driving patterns of community assem-bly and disassembly is therefore critical to linking com-munity dynamics to ecosystem properties (Weiher andKeddy 1999b, Thompson et al. 2001, Dıaz et al. 2003).

In addition, more effort is needed to reconcile con-trasting predictions from cross-system comparisonsand local-scale experimental results. For example,while several modeling and experimental studies in-dicate complementarity among plants (see SectionII.A.2, above), Grime (2001) predicts that comple-mentarity will primarily occur where species richnessand productivity are positively correlated on the risingpart of the unimodal diversity/productivity curve. Heargues that effects of legumes on nitrogen availabilitycan skew results if experimental plots do not have timefor species richness to equilibrate to the higher fertilityinduced by N-fixation. These are testable hypothesesthat could help resolve the ongoing debate about mech-anisms of diversity effects on productivity (see SectionII.C.2, below). On another front, allometric scaling re-lationships (e.g., resource use vs. organism size) arestrong across many orders of magnitude, despite ig-noring functional differences among species other than

size (e.g., Enquist et al. 1998). The overall pattern sug-gests little effect of functional differentiation or speciesdiversity on ecosystem properties. However, substan-tial variation (10–100 fold) in the size–resource useregression exists within the scales relevant to localcommunity studies, implying that community interac-tions, species’ functional differences (other than size),and diversity can be ecologically quite significant atsuch scales (Chapin et al. 1996b, Lavorel and Garnier2002, Mittelbach et al. 2003, Whittaker and Heegaard2003). Ecologists need to investigate just how strongthe effects of diversity are relative to other ecosystemcontrols and how such relationships change across eco-logical scales (see also Section I.C., above).

2. Matching inference with experimental design.—The debate about interpretation of experiments (e.g.,Huston 1997, Wardle 1999, Huston et al. 2000, Wardleet al. 2000b, Huston and McBride 2002) emphasizesthe point that care is needed regarding inferences inboth observational and experimental studies about (1)effects of biodiversity relative to abiotic controls and(2) the application of results to real world scenarios ofspecies extinctions. The controversy over experimentscomes from two criticisms. First, selective reductionof diversity in response to underlying variables suchas fertilization leads to the problem of ‘‘hidden treat-ments’’ (Huston 1997), which in turn leads to difficultydetermining the mechanisms underlying changes inecosystem response. Experiments incorporating ran-domized–combinatorial species assemblages became apopular experimental design in response to this prob-lem (Naeem et al. 1995, Hector et al. 1999, Tilman1999, Schmid et al. 2002a). These experiments attemptto understand the functional consequences of speciesloss from a regional pool (Vitousek and Hooper 1993,Lawton 1994, Naeem et al. 1995, Sala et al. 1996), forwhich random, synthetic communities provide onemeans of exploring the space of ‘‘possible communi-ties.’’

The primary issue is how to interpret such approach-es in the context of actual scenarios of biodiversitychange, which is the second criticism leveled at syn-thetic community studies. The randomized-assemblageexperimental design may not resemble the way realcommunities are assembled (or disassembled, in thecase of extinctions; Wardle et al. 2000b) and it createsa complex experiment with multiple autocorrelations(Huston and McBride 2002, Naeem 2002a). Such dif-ficulties, in combination with the hidden treatmentsissue, mean that experimental designs must be closelymatched to the primary aims and hypotheses of eachstudy (Schmid et al. 2002a).

The relevance of any such experiments to the realworld depends on the drivers of community change andthe patterns of species loss or introduction. Species


losses may or may not be random with respect to spe-cies’ effects on ecosystem properties, depending onwhether traits related to response to the environmentalso affect ecosystem properties (Grime 1979, Chapin1980, Fridley 2001, Lavorel and Garnier 2002). Insome studies, for example, the best predictor variablesfor plant extinction are life-history characteristics suchas low seed dormancy, short life cycle, and absence ofclonality, which are often independent of functionaltraits influencing ecosystem properties (e.g., Chapin etal. 1996a, Grime et al. 1997a, Stocklin and Fischer1999). In such cases, randomized combinatorial ex-periments may provide a reasonable approximation toactual extinction scenarios (Loreau et al. 2001).

On the other hand, losses in plant and animal di-versity are often not random and often do not occuragainst an unchanging abiotic background. Species canbe lost from particular sites by the same processes thatare destroying or transforming the ecosystems that con-tain them, resulting in functional shifts in the biota.That is, sets of species with particular functional traitsare being replaced by other sets with different traits(Ratcliffe 1984, Thompson 1994, Janssens et al. 1998).In such cases, random scenarios could be misleadingbecause differences in extinction probability amongspecies may be driven by species’ traits directly linkedto effects on processes (Duncan and Young 2000). Forexample, nutrient enrichment consistently selects forfast-growing species that outcompete slower-growing,more stress-tolerant species (Tilman 1987, Aerts et al.1990, Wedin and Tilman 1993, MacGillivray et al.1995). Given this consistency of diversity and domi-nant species’ traits across gradients of nitrogen depo-sition, the debate about effects of species richness vs.composition in some experiments (Givnish 1994, Til-man and Downing 1994, Huston 1997) may be mootfrom the perspective of managing nitrogen within land-scapes. Additional research is needed to understandhow the functional traits of species interact with dif-ferent types of human-caused environmental changesto determine pathways of species loss (Sala et al. 2000,Chapin et al. 2001). How might these pathways of spe-cies loss influence various ecosystem properties andservices? How might these changes feed back to furtherbiodiversity changes?

Using a variety of experimental approaches will con-tribute to robust answers to these questions. Removalexperiments offer potential to address such questionsbecause they directly explore what happens to an eco-system when a species or functional type is no longerpresent (e.g., Sala et al. 1989, Hobbie et al. 1999, War-dle et al. 1999, Symstad and Tilman 2001, Dıaz et al.2003). Of course, manipulative experiments, using ei-ther synthetic communities or species removals, haveother constraints, such as being restricted to easily ma-

nipulated ecosystem types, time needed for establish-ment or disturbance recovery, and other issues asso-ciated with scale and realism such as being too small,too brief, or too simple in structure when compared tonatural systems. For example, natural patterns of spa-tial aggregation of species can be an important factoraffecting both species coexistence and, potentially,ecosystem properties (e.g., Freckleton and Watkinson2000, Stoll and Prati 2001). For these reasons, there isan urgent need for observational studies carefully de-signed to determine whether the patterns and mecha-nisms predicted by theories and experiments occur innatural systems (Troumbis and Memtsas 2000, but seealso Wardle 2001b).

III. MANAGEMENT IMPLICATIONS

Biodiversity, whether at the level of genes, species,or communities, clearly affects the way ecosystemsfunction, as outlined in the preceding sections. Theprotection of ecosystem goods and services alreadyforms an important part of environmental theory andpractice (Daily 1997). Wildlife and habitat conserva-tion, the Convention on Biodiversity, intergovernmen-tal panels and agencies, and international scientific pro-grams represent widespread activities designed to un-derstand and reduce biodiversity loss and species in-vasions. Incorporating what we know about the basicroles that biodiversity can play in ecosystem propertiescan help to inform the development of more effectiveenvironmental management and policy, and improveour abilities to predict environmental change. Yet, ifdiversity is so often crucial to sustained functioning ofecosystems, why is it not always a routine part of man-agement philosophy for farmers, ranchers, fishers, andforesters?

In this section, we seek to address this question, andto suggest areas in which application of the principlesdiscussed in Section II might be most beneficially ap-plied. Our aim is to be suggestive, rather than exhaus-tive, in coverage. The needs for knowledge on biodi-versity’s effects on ecosystem properties and servicesclearly depend on the type of management questionsbeing addressed. These questions cover a huge rangeof activities and scales, from small-scale, local issues(e.g., how many and which species might be necessaryto produce food and reduce erosion for a farmer in thetropics), to regional issues (e.g., how to manage forestlands for wood production, fish and wildlife habitat,and recreation), to global issues (e.g., how might shiftsin species composition and diversity associated withclimate change influence carbon sequestration in thebiosphere). We discuss a few key examples in hopesof illustrating potential applications. Here, we differ-entiate managed ecosystems from merely degradedecosystems (‘‘polluted, overexploited, or converted


FIG. 6. Anticipated effects of diversity on ecosystem properties (plant net primary productivity is shown) across increasingscales. As habitat heterogeneity, temporal variation in conditions, and response to disturbance are included, more speciesare needed to saturate ecosystem properties. If species are selected at random, rather than chosen according to their adaptations,ecosystem properties may saturate even more slowly. ‘‘Zone accessible to intensive management’’ reflects agronomic eco-systems where very high productivity may be achieved at very low species richness, but at the cost of substantial inputs oftime, energy, fertilizers, pesticides, and/or water resources, often with concurrent off-site impacts and trade-offs with otherecosystem services. The figure is modified from Field (1995).

that have not been managed’’; cf. Silver et al. 2001).While systems that are intensively managed for pro-duction differ in many respects from lightly managednatural and seminatural ecosystems, application of theprinciples discussed in Section II is similar in somecritical situations, particularly those involving increas-es of scale.

In general, management of production-oriented sys-tems has focused on composition (the species or ge-notypes that perform best under a given set of condi-tions) rather than species diversity. Lack of diversityin intensively managed systems occurs for a simplereason: diversity, whether of genotypes, species, orcommunities, often complicates management, makingit less profitable in the short term (Ewel 1991). Espe-cially with industrialized agriculture, diversity posesproblems for automated planting and harvesting andefficiencies of scale, and this is not a cost that manyenterprises are willing to incur. The ecological per-spectives of more complete resource utilization or abil-ity to sustain natural levels of biodiversity are seldomincorporated into the decision-making process, eitherat the levels of individuals or government policy. Uni-formity has been the rule, and it tends to be driven byeconomic, not ecological, considerations (Ewel 1991).

Given the apparent conflict between near-term profitsand longer term sustainability, are we inevitably headedfor a world whose oceans are managed for a handfulof pelagic fishes and whose lands are a simplified patch-work of soybeans, cows, wheat, rice, pines, eucalypts,

corn, concrete, and cosmopolitan weeds? No, as longas broader scale factors are incorporated into calcula-tions of management costs. Maintenance of high pro-ductivity over time in monocultures almost invariablyrequires heavy subsidies of chemicals, energy, and cap-ital (Fig. 6; Field 1995), and it may not be possible orsustainable in some systems in the face of various dis-turbances, diseases, soil erosion, overuse of naturalcapital (e.g., water), or trade-offs with other ecosystemservices (Ewel 1999, Postel 1999). Diversity becomesincreasingly important as a management goal, fromboth economic and ecological perspectives, with in-creasing temporal and spatial scales and for providinga broader array of ecosystem services.

Diversity is consciously incorporated into ecosys-tems managed for extraction of food and fiber underat least three circumstances: (1) as a safeguard againstrisk resulting from, for example, fluctuations in envi-ronmental conditions or marketplace demands; (2) asa means of extending use of a site’s resources overtime; and (3) as a means to provide multiple goods andservices. A variety of practices illustrate these excep-tions to ‘‘management by imposed homogeneity’’ inintensively managed systems, ranging from multiple-use forests and rangelands, to intercropping and ag-roforestry, to the home gardens so common in manytropical countries. For example, home gardens providea variety of food, medicinal, and construction materials(e.g., Watson and Eyzaguirre 2002), and public man-agement of forests provides for a variety of goods and


services, including many that are non-extractive. Ge-netic diversity of crops decreases susceptibility to pestsand climate variation (Ewel 1986, Altieri 1990, Zhu etal. 2000). Especially in low-input systems, locallyadapted varieties often produce higher yield or are moreresistant to pests than varieties bred for high perfor-mance under optimal conditions (Joshi et al. 2001).Intercropping, crop rotations, and mixed-species for-estry often include legumes to increase soil nutrientcapital (Trenbath 1974, Vandermeer 1989, Cannell etal. 1992). Diversity of pasture species can reduce nu-trient leaching, production variation, and insurancecosts (Schlapfer and Erikson 2001, Mader and Fliess-bach 2002, Schlapfer et al. 2002). In all of these cases,more knowledge-based, biologically detailed manage-ment may require greater short-term effort or expense,but with the benefit of longer term sustainability and/or reduced off-site impacts.

Increasing the temporal scale of management is rel-evant in less intensively managed systems as well. Forexample, concern about carbon dioxide emissions intothe atmosphere has generated substantial discussionabout reducing the net impact of those emissionsthrough biotic sequestration. If it is to be effective,carbon sequestration must be a long-term undertaking,involving sustainable rates of fixation and storage.High rates of net primary production (NPP) that onemight witness from a monoculture in the short term arelikely to drop with time, or at least oscillate consid-erably, as pest attacks, cohort senescence, and weathervariability take their tolls. As discussed in SectionII.B., diversity can provide compensatory capabilityrequired over time if an ecosystem is to sustain C se-questration (e.g., Bolker et al. 1995), including re-sponses of NPP to elevated CO2 and nitrogen (Stockeret al. 1999, Niklaus et al. 2001b, Reich et al. 2001, Heet al. 2002). NPP is only half the story, however. Netecosystem production (NEP; total C uptake minus allC losses, both autotrophic and heterotrophic) is thecritical measure for C sequestration (Catovsky et al.2002). We have considerably less information aboutplant diversity effects on the suite of processes influ-encing C losses from ecosystems, such as decompo-sition and heterotrophic respiration, than for NPP (seeSection II.A.2). Additional research on whole systemC cycling in response to varying plant diversity is acritical need for the future (Catovsky et al. 2002, Diaset al. 2003).

The need for incorporating diversity also arises whenconsidering management over large spatial scales andfor multiple ecosystem services. The importance offunctional diversity at the ecosystem or landscape scaleis well known in a number of specific cases. For ex-ample, areas of natural habitat can serve as sources ofpropagules for recolonization of sites affected by other

stresses (Cushman et al. 1995). Considering the spe-cies–area relationship, higher species diversity at thelandscape scale is necessary to maintain even moderatediversity at the local scale (Wilson and Willis 1975).Connectedness among a variety of ecosystem types isnecessary for some far-ranging species such as migra-tory birds and fish, and for predators with alternateprey, some of which may be agricultural pests (Cush-man et al. 1995, Perfecto et al. 1996, Greenberg et al.1997). In all these cases, effects may be mediated bya diversity of species or by particularly strong effectsof a given species or functional type (Noble and Burke1995). For parks and preserves managed for biodiver-sity preservation, potential effects of climate changeon species’ ranges necessitate managing diversity atthe landscape to regional scale. As climate patternschange, the organisms best adapted to a particular suiteof environmental conditions may find themselves un-able to migrate from recently changed conditions (e.g.,Etterson and Shaw 2001). The need for diversity atlarge spatial scales is already being implemented in theconservation community (e.g., see the Conservation byDesign Initiative of The Nature Conservancy and theGlobal 200 Ecoregions of the World Wildlife Fund),but it is not yet widely recognized by those engagedin management for food and fiber.

In intensively managed systems, management choic-es influence the diversity of associated species as wellas those species targeted for harvest, both within theharvesting area and in the surrounding landscape (Fig.7; Fischer and Stocklin 1997, Giller et al. 1997, Gon-zalez et al. 1998, Vandermeer et al. 1998, 2002, Stock-lin and Fischer 1999, Vandermeer and Carvajal 2001).Changes in composition and diversity of associatedspecies have potential consequences for forest or ag-ricultural production via supporting ecosystem services(e.g., generation of soil fertility, pest control, polli-nation), as well as for additional services (e.g., pro-vision of wildlife habitat) (Pimentel et al. 1992, Nabhanand Buchmann 1997, Naylor and Ehrlich 1997). Thereare many management strategies for increasing diver-sity of associated species (e.g., Hanson et al. 1991,Pimentel et al. 1992), which is often important for bio-diversity value itself. However, the net effects on pro-duction of target species are variable, depending on theparticular functional relationships among the speciesinvolved, as discussed above (Section II.B.2(b); Ewel1986, Vandermeer et al. 2002). Effects of diversity onpest and pathogens are not restricted to agriculturalsituations. In deciduous forests of eastern North Amer-ica, shifts in trophic relations among mammals canaffect the degree to which humans are exposed to Lymedisease and other pathogens (Ostfeld and Keesing2000). Theory predicts that under some conditionsmammalian diversity is an important component of the


FIG. 7. Potential patterns of effects of intensification ofagricultural practices on diversity of nontarget species. Let-ters a–f on the x-axis refer to increasing states of managementintensity, with ‘‘a’’ being an unmanaged ecosystem and ‘‘f’’being intensive, industrialized agriculture. Intensificationtends to reduce diversity of associated taxa, although the pat-terns could follow a variety of trajectories, including the po-tential for initial increases in species richness for some taxaunder the assumptions of the intermediate disturbance hy-pothesis (Giller et al. 1997). Losses of associated diversitymay thereby affect ecosystem services related to agriculturalproduction, although the effects often depend on the detailsof the relationships among the species and services in ques-tion (see Section III). The figure is modified from Vandermeeret al. (2002).

response, but under other conditions, the functionaltraits of certain key species have a dominant effect ondisease prevalence (Schmid and Ostfeld 2001).

Management for multiple goals or functions becomesmore complicated because of potential trade-offsamong them. At what points (e.g., across a gradient ofintensified land use) might species be lost whose effectson services either cannot be replaced, or are very ex-pensive to replace? What are the ecological and eco-nomic trade-offs associated with these losses (e.g.,Balmford et al. 2002)? Understanding which species(both target and associated) affect which ecosystemproperties and services will help in this regard, but itmay take many years to understand fully. In the interim,adaptive management and maintaining a diversity ofnative species will help maintain future managementoptions.

Biodiversity, broadly defined, clearly has strong ef-fects on ecosystem properties and the goods and ser-vices derived from them. Similarly, there is no debateabout the effects of the functional characteristics ofparticular species. Most controversy has surrounded theeffects of species richness, which are more variable.Yet, despite some mechanistic uncertainties and the

economic complexities caused by incorporating ge-netic, species, and community diversity into manage-ment, there is little doubt that this needs to be done,especially in making decisions involving large tem-poral and spatial scales. Sacrificing those aspects ofecosystems that are difficult or impossible to recon-struct, such as diversity, simply because we are not yetcertain about the extent and mechanisms by which theyaffect ecosystem properties, will restrict future man-agement options even further (Costanza et al. 1998,Lauck et al. 1998). It is incumbent upon ecologists tocommunicate this need, and the values that can derivefrom such a perspective, to those charged with eco-nomic and policy decision-making.

SUMMARY

Human-induced changes to components of theEarth’s biodiversity have the potential to compromisethe performance of ecosystems, both immediately andby impeding their ability to respond to altered condi-tions. A long history of ecological experimentation andtheory supports the postulate that ecosystem goods andservices, and the ecosystem properties from which theyare derived, depend on biodiversity, broadly defined.Functional traits of species are important drivers ofecosystem properties, and we are learning more abouthow these traits combine to affect properties in morediverse systems. While the debate surrounding someexperiments on the effects of plant species richness onecosystem properties has received some negative press,in the end it has helped broaden and deepen ecologicalunderstanding and point the way for future studies in-tegrating both community and ecosystem perspectives(Loreau et al. 2001, 2002b, Kinzig et al. 2002, Naeem2002b). Species composition, richness, evenness, andinteractions all both respond to and influence ecosys-tem properties. Further progress will require integra-tion of knowledge about how communities are struc-tured with knowledge about controls on ecosystemproperties. To make links to ecosystem management,further understanding of the social and economic con-straints of potential practices needs to be integratedwith our ecological knowledge (e.g., the MilleniumEcosystem Assessment, available online).18 The mech-anisms of biodiversity effects are likely to differ amongecosystem properties, ecosystem types, managementgoals, and pathways of potential biodiversity changes.Understanding this complexity, while taking strongsteps to minimize current losses and invasions of spe-cies, is an important step toward our ability to respon-sibly manage Earth’s ecosystems and the diverse biotathey contain.

18 ^http://www.millenniumassessment.org/en/index.aspx&


ACKNOWLEDGMENTS

Peter Vitousek contributed greatly to the committee pro-cess. Phil Grime, Michael Huston, Christa Mulder, Peter Vi-tousek, and an anonymous reviewer provided extensive com-ments on earlier drafts of the manuscript. Phil Grime alsoparticipated in the initial meeting and formulation of thisdocument. Tim Kreps gathered and helped interpret literature.Katherine Cottingham, Peter Morin, and David Tilman con-tributed figures. The initial committee meeting and ideas forthe paper benefited greatly from the conference ‘‘Biodiversityand ecosystem functioning: synthesis and perspectives,’’ 6–9 December 2000, organized by Michel Loreau, ShahidNaeem, and Pablo Inchausti under the auspices of IGBP-GCTE Focus 4 and DIVERSITAS Core Programme Element1, and funded by the European Science Foundation, the CentreNational de la Recherche Scientifique (France), and the Na-tional Science Foundation (USA). Anonymous donors to theEcological Society of America provided additional meetingsupport. Partial support for D. Hooper came from NSF grantsDEB-9974159 and DEB-0213187.

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